Restoration of invaded Cape Floristic Region riparian

ORIGINAL PAPER

Restoration of invaded Cape Floristic Region riparian systemsleads to a recovery in foliage-active arthropod alpha- andbeta-diversity

Malebajoa A. Maoela1,3 • Francois Roets1,3 • Shayne M. Jacobs1,4 •

Karen J. Esler1,2,4

Received: 28 August 2015 / Accepted: 29 December 2015 / Published online: 12 January 2016

� Springer International Publishing Switzerland 2016

Abstract The Cape Floristic Region of South Africa is a

global biodiversity hotspot threatened by invasive alien

plants (IAPs). We assessed the effect of plant invasions,

and their subsequent clearing, on riparian arthropod

diversity. Foliage-active arthropod communities were col-

lected from two native and one invasive alien tree species.

Alpha- and beta-diversity of their associated arthropod

communities were compared between near pristine, Aca-

cia-invaded and restored sites. Arthropod alpha-diversity at

near pristine sites was higher than at restored sites, and was

lowest at invaded sites. This was true for most arthropod

taxonomic groups associated with all native tree species

and suggests a general trend towards recovery in arthropod

alpha-diversity after IAP removal. Overall, arthropod spe-

cies turnover among sites was significantly influenced by

plant invasions with communities at near pristine sites

having higher turnover than those at restored and invaded

sites. This pattern was not evident at the level of individual

tree species. Although arthropod community composition

was significantly influenced by plant invasions, only a few

significant differences in arthropod community composi-

tion could be detected between restored and near pristine

sites for all tree species and arthropod taxonomic groups.

Assemblage composition on each tree species generally

differed between sites with similar degrees of plant inva-

sion indicating a strong turnover of arthropod communities

across the landscape. Results further suggest that both

arthropod alpha- and beta-diversity can recover after IAP

removal, given sufficient time, but catchment signatures

must be acknowledged when monitoring restoration

recovery.

Keywords Acacia mearnsii � Riparian zone � Invasive

alien plants � Arthropod responses

Introduction

Terrestrial arthropod populations and communities are

associated with certain vegetation types, and the loss of

suitable plant habitat can lead to their declines (Herrera and

Dudley 2003; Longcore 2003). Among the primary threats

to arthropod diversity are introductions of invasive species

(Tallamy 2004; Magoba and Samways 2012). Dense stands

of invasive alien plants (IAPs) are a growing threat to

native biodiversity and ecosystem functioning (Sala et al.

2000; Le Maitre et al. 2004; Clavero and Garcia-Berthou

2005; van Wilgen et al. 2008). They cause changes in

vegetation structure, composition and host quality and

therefore affect arthropod assemblages (Beerling and

Dawah 1993). For example, Slobodchikoff and Doven

(1977) showed that increased cover of the non-native grass

Ammophilia arenaria disrupted the structure of sand dune

arthropod communities in California. Similarly, abundance

and composition in native ant and bird communities has

& Malebajoa A. Maoela

[email protected]

1 Department of Conservation Ecology and Entomology,

Stellenbosch University, Private Bag X1, Matieland 7602,

South Africa

2 DST/NRF Centre of Excellence for Invasion Biology,

Stellenbosch University, Private Bag X1, Matieland 7602,

South Africa

3 DST/NRF Centre of Excellence in Tree Health

Biotechnology (CTHB), Forestry and Agricultural

Biotechnology Institute (FABI), University of Pretoria,

Private Bag X20, Hatfield, Pretoria 0028, South Africa

4 Water Institute, Stellenbosch University,

Private Bag X1, Stellenbosch 7602, South Africa

123

J Insect Conserv (2016) 20:85–97

DOI 10.1007/s10841-015-9842-x

http://orcid.org/0000-0001-5886-7991

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been altered by IAPs in the Cape Floristic Region (CFR) of

South Africa (French and Major 2001; Mokotjomela and

Hoffmann 2013), with significant implications for the seed

dispersal of native plants (Mokotjomela and Hoffmann

2013).

We investigate the effect of invasive alien trees on

arthropod assemblages associated with native riparian trees

in the CFR, a region heavily impacted by woody IAPs.

Riparian ecosystems are among the most endangered CFR

habitats, with \20 % of their original extent still intact

(Nel et al. 2007). Riparian vegetation is used for resting,

feeding, reproduction and refuge by both aquatic and ter-

restrial arthropods, and provides a critical resource base for

vertebrates (Gray 1993). One of the most notorious inva-

sive species in the CFR is Acacia mearnsii, which the

Working for Water (WfW) invasive plant clearing pro-

gramme has designated as a top priority for removal (van

Wilgen et al. 2008, 2012). Most South African research on

A. mearnsii and other IAPs in the Fynbos biome has shown

that dense stands of invasive acacias can rapidly reduce the

abundance and diversity of native plants at the landscape

scale (Richardson et al. 1989). Dense stands of IAPs also

lead to a decline in soil seed banks of riparian systems

(Vosse et al. 2008), increasing the probability of extinc-

tions of native species. In addition, IAPs greatly increase

biomass (Milton 1981), affect fire regimes (Van Wilgen

et al. 2008), change nutrient cycles (Witkowski 1991) and

reduce arthropod richness (reviewed by Litt et al. 2014).

Clearing of IAPs can lead to recovery of vegetation

communities under certain conditions (Blanchard and

Holmes 2008) and it can be expected that removal of A.

mearnsii from riparian systems would also help restore the

high arthropod species diversity that characterizes CFR

riparian communities (Samways et al. 2011). Studies have

shown that arthropod richness (alpha-diversity) and abun-

dance can recover after restoration efforts on disturbed

riparian ecosystems (Williams 1993; Longcore 2003;

McCall and Pennings 2012). Removal of invasive Phrag-

mites resulted in the return of dominant native vegetation

and the re-establishment of arthropod species assemblages

(Gratton and Denno 2005). The effect of clearing of IAPs

from CFR riparian ecosystems on riparian arthropod

diversity has not yet been assessed.

Although the advantages of the removal of IAPs are

apparent, the process itself represents yet another distur-

bance to river ecosystems. IAP clearing can result in

unexpected changes to ecosystem processes that may affect

arthropod survival. For example, removal of IAPs alters

canopy characteristics, which directly affects the interior

environments of ecosystems (i.e., temperature, humidity,

and radiation), this, in turn, leads to changes in arthropod

richness and abundance (Ziesche and Roth 2008). Apart

from microclimate, altered architectural habitat complexity

(Schowalter and Crossley 1988) and changes in plant

nutritional quality (Fischer et al. 2010), IAP removal may

also influence arthropod communities by limiting their

dispersal ability by creating isolated patches (Schowalter

and Crossley 1988).

The measurement of arthropod species richness (alpha-

diversity) and species turnover (beta-diversity) under dif-

ferent levels of plant invasions aids our understanding of

the effect of management conservation of these systems

(Kessler et al. 2009). We test the effect of an invasive alien

tree on arthropod alpha- and beta-diversity in riparian

ecosystems of the CFR and whether measures of arthropod

alpha- and beta-diversity can indicate a trajectory of

recovery post-IAP removal. We expected to see differences

in arthropod alpha- and beta-diversity among areas differ-

ing in degree of invasion (near pristine, heavily invaded by

A. mearnsii and cleared ca. 7 years prior to the com-

mencement of this study), with major differences between

the near pristine habitats and those that have been restored

reflecting the probable time it takes for arthropod assem-

blages to fully recover after invasion and subsequent

mechanical clearing of IAPs.

Materials and methods

Study area and species

This study was conducted in the mountain stream and

foothill sections of several riparian systems within the

Western Cape, South Africa (Fig. 1; Table 1). The selected

river reaches are on quartzitic sandstone that is character-

istically acidic and low in nutrients and dissolved solids

(Day and King 1995; Table 1). Vegetation is largely

shrubby Fynbos and includes a variety of tree taxa that

form forest patches (Goldblatt and Manning 2000).

Nine study sites in the three different rivers systems

were identified: three near pristine sites (NP) (reference

sites), three heavily invaded (HI) sites (predominantly by

A. mearnsii), and three restored sites (R) (formerly invaded

sites that had been cleared of IAPs more than 7 years prior

to this study). Site categorisation into near ‘pristine’ or

‘heavily invaded’ was based on visual scoring of Acacia

mearnsii cover within two transects measuring 50 m in

length (parallel to the river) and 5 m in width (perpendic-

ular to the river and crossing both wet and dry bank zones).

For heavily invaded sites A. mearnsii canopy cover was

[75 % and for near pristine was\5 %. For restored sites,

site categorisation was based on the IAPs clearing history

of the sites. In restored sites, IAPs were felled as close to

the base as possible and herbicide was applied to stumps.

Potential sites were identified using information obtained

from previous studies (Blanchard and Holmes 2008) and

86 J Insect Conserv (2016) 20:85–97

123

confirmed by discussions with conservation authority

managers (CapeNature), members of WfW and private

landowners.

For the purposes of this study, two tree species endemic

to Fynbos riparian zones, Brabejum stellatifolium (L.)

(Proteaceae) and Metrosideros angustifolia (L.) (Myr-

taceae) were selected. These trees are naturally confined to

the Fynbos (Thuiller et al. 2006) where they prefer moist

areas and therefore commonly occur along streams

(Mucina and Rutherford 2006). These tree species are

abundant and important components of riparian habitats

and considered key species in south-western Cape

Mediterranean-type riparian systems (Galatowitsch and

Richardson 2005). In addition to these two native species,

the woody invasive alien species Acacia mearnsii DeWild

(L.) (Fabaceae) that commonly invades habitats dominated

by B. stellatifolium and M. angustifolia was selected.

Acacia mearnsii was chosen because it commonly invades

habitats dominated by B. stellatifolium and M. angustifolia

and also, its seed germination is usually prompted by

disturbances.

Arthropod collection

As CFR arthropods show substantial seasonal variation

(Roets and Pryke 2013), sampling was conducted once

during summer (2011), autumn (2011), winter (2012), and

spring (2012) and the data from all four seasons were

combined for analyses. The sampling was done within two

transects measuring 50 m in length (parallel to the river)

and 5 m in width (perpendicular to the river and crossing

both wet and dry bank zones). Arthropods associated with

the foliage of the three tree species were sampled using a

petrol-driven Blow and Vac (Stihl, Germany) suction

apparatus (Stewart and Wright 1995). Five individuals of

each of the three tree taxa of similar height and stem

diameter were selected at random at each site and arthro-

pods collected from their crowns by inserting tips of

branches into the nozzle for 30 s. This process was repe-

ated 70 times on different branches for each individual tree.

Catches per individual tree were kept separate. Collected

arthropods were transferred to re-sealable plastic bags,

stored at -20 �C, and later assigned to morphospecies and

taxonomic order (Oliver and Beattie 1996). Reference

material was stored in 70 % ethanol and is held at the

University Stellenbosch Insect Collection (USEC), Stel-

lenbosch, South Africa.

Statistical analyses

A non-parametric richness estimator was selected, to

establish sampling representativity because most arthropod

assemblages normally have a large number of rare species

(Hortal et al. 2006). The Chao2 estimator was used as it is

considered to be the least biased and most precise estimator

when working with small sample sizes (Walther and

Morand 1998). Values were calculated using EstimateS

(Colwell 2009).

Fig. 1 Location of the three

Western Cape rivers: 1 Dwars, 2

Molenaars, and 3 Wit and the

nine sites (circle near pristine,

square heavily invaded and

triangle restored) used in this

study

J Insect Conserv (2016) 20:85–97 87

123

Table

1S

ite-

spec

ific

info

rmat

ion

,in

clu

din

gth

em

ajo

rg

eom

orp

ho

log

ical

char

acte

rist

ics,

site

trea

tmen

tan

dm

ean

ann

ual

rain

fall

Sit

eG

eolo

gy

Tre

atm

ent

of

inv

asio

n

His

tory

clea

ran

ceF

ire

his

tory

Mea

nan

nu

al

rain

fall

(mm

)

Lo

ng

itu

din

alzo

ne

Co

ord

inat

es

Nearpristine

Up

per

Dw

ars

(UD

)S

and

sto

ne/

gra

nit

eN

on

eN

on

eN

op

roo

fo

fla

test

fire

57

8M

ou

nta

inh

ead

wat

er

stre

am

33�5

70 0

5.5

400 S

;

18�5

80 3

9.2

200 E

Elevation

44

6m

Bai

ns

Klo

of

(BK

)S

and

sto

ne

No

ne

No

ne

No

pro

of

of

late

stfi

re8

88

Mo

un

tain

hea

dw

ater

stre

am

33�3

40 0

8.4

900 S

;

19�0

80 1

9.0

300 E

Elevation

30

3m

Du

toit

sK

loo

f(D

K)

San

dst

on

eN

on

eN

on

eN

op

roo

fo

fla

test

fire

14

68

Fo

oth

ill

33�4

30 4

7.4

100 S

;

19�0

60 3

7.0

600 E

Elevation

47

2m

Heavily

invaded

Lo

wer

Wit

(LW

)S

and

sto

ne

No

ne

No

ne

20

12

83

3F

oo

thil

l3

3�3

20 1

9.5

600 S

;

19�1

00 5

1.7

700 E

Elevation

24

3m

Mid

Wit

(MW

)S

and

sto

ne

No

ne

No

clea

rev

iden

ce2

01

28

33

Fo

oth

ill

33�3

40 0

6.3

300 S

;

19�0

80 4

7.5

200 E

Elevation

28

3m

Mid

Dw

ars

(MD

)S

and

sto

ne/

gra

nit

eN

on

eN

ocl

ear

evid

ence

No

pro

of

of

late

stfi

re5

78

Mo

un

tain

stre

am3

3�5

60 5

3.3

600 S

;

19�5

80 1

1.2

500 E

Elevation

40

0m

Restored

Up

per

Mo

len

aars

(UM

)S

and

sto

ne

[7

yea

rsag

o(A.

mearnsii)

Init

ial

trea

tmen

t:

20

02

–2

00

3

2fo

llo

w-u

p

trea

tmen

ts.

(fel

l

and

rem

ov

e)

20

12

88

9U

pp

erfo

oth

ill

33�4

20 3

8.5

600 S

;

19�1

10 4

9.2

400 E

Elevation

33

5m

Du

To

it(D

T)

San

dst

on

e[

7y

ears

ago

(A.

mearnsii)

Init

ial

trea

tmen

t:

20

02

(fel

lan

d

rem

ov

e)

No

pro

of

of

rece

nt

fire

14

77

Up

per

foo

thil

l3

3�4

30 3

4.2

100 S

;

19�0

60 0

1.0

200 E

Elevation

54

4m

Lo

wer

Dw

ars

(LD

)S

and

sto

ne

[8

yea

rsag

o(A.

mearnsii

andA.

longifolia

)

Init

ial

trea

tmen

t:

20

02

.3

foll

ow

-up

trea

tmen

ts(f

ell

and

rem

ov

e)

No

pro

of

of

rece

nt

fire

.5

78

Mo

un

tain

stre

am3

3�5

60 4

5.7

400 S

;

18�5

70 5

7.5

100 E

Elevation

38

5m

All

the

stre

ams

are

per

enn

ial


123

Arthropod alpha-diversity (a) (or species richness) for

heavily invaded, restored and near pristine riparian plant

invasions was compared using generalised linear models

(GLMs). These variables were fitted to a Poisson distri-

bution model with a log-link function using generalised

estimating equations (Zuur et al. 2010) in Proc Genmod of

SAS 9.1 (SAS Institute Inc., Cary, USA). The Poisson

distribution type was selected to minimize the deviance

statistic (Johnson et al. 2006). Test statistics were calcu-

lated using the penalised quasi-likelihood technique, as

variances showed no over-dispersion (Bolker et al. 2008).

Separate analyses were run for the three host tree species,

sites within each plant invasions type, as well as for the

eight most species rich arthropod taxonomic groups (Ara-

neae, Coleoptera, Diptera, Hemiptera, Hymenoptera,

Lepidoptera, Ants, and Orthoptera). Significant differences

under this model are reported where p B 0.05.

Two measures of beta-diversity were assessed in this

study: (1) b1 = species turnover among sites of the same

plant invasions (Anderson 2006) and (2) b2 = assemblage

compositional changes between sites with different plant

invasions (Anderson 2006; Pryke et al. 2013). Species

turnover among sites of the same plant invasions (b1) was

calculated using a resemblance matrix based on the Jaccard

measure. The Jaccard dissimilarity measure uses only

compositional (presence/absence) information and is

directly interpretable as the percentage of unshared species

among samples (Terlizzi et al. 2009). To determine the

variability in species composition within the study sites, the

Permutational Analysis of Multivariate Dispersions

(PERMDISP) routine in the Permutational Multivariate

Analysis of Variance (PERMANOVA?) extension in

PRIMER 6 was conducted. PERMDISP (b1-diversity)

determines the mean distance of samples to the geometric

centre (centroid) of each predefined group (e.g. arthropods

associated with A. mearnsii from near pristine sites) in

three dimensional space (Anderson 2006). This allows for

comparisons between the mean distances to various cen-

troids (e.g. arthropods associated with A. mearnsii from

near pristine, restored and heavily invaded sites respec-

tively) using ANOVA to determine F- and p values (An-

derson 2006) and allows for pair-wise testing. These

analyses were performed in PRIMER 6 (PRIMER-E 2008)

with 9999 permutations (Anderson 2006).

Compositional differences across different plant inva-

sion status (near pristine, heavily invaded and restored) and

sites within each plant invasion status (b2) (Anderson

2006) were compared using PERMANOVA? in PRIMER

6. The F and p values for the main test (as well as t values

for pair-wise differences) for similarity of the eight taxo-

nomic groups listed above between each plant invasion

type and the three host trees were calculated using 9999

permutations. Hierarchical agglomerative clustering

analyses were performed using Bray–Curtis similarity

(Bray and Curtis 1957) after fourth-root transformation of

data to reduce the influence of common species (Anderson

2001). Results were visually represented using Principal

Coordinates Ordination (PCO) plots (Clarke 1993) in

PRIMER 6. Diversity indices were compared for all tree

taxa combined, for each individual tree species and for the

eight most species rich arthropod taxonomic groups.

Results

Arthropod alpha-diversity

A total of 29,811 arthropod individuals representing 967

morphospecies from 15 orders were collected. The most

abundant orders were the Coleoptera (14,253), Hemiptera

(5197), Diptera (3359), Araneae (1734), Hymenoptera

(1710) (excluding the Ants, 470), Lepidoptera (237), and

Orthoptera (388). The near pristine sites had the highest

number of observed and estimated species, while the

heavily invaded sites had the lowest number of observed

and estimated species (Table 2). The restored sites had

intermediate numbers for observed and estimated number

of species (Table 2). For M. angustifolia, the estimated

numbers of species at heavily invaded sites were similar to

those at restored sites (Table 2). For B. stellatifolium and A.

mearnsii the estimated number of species varied across all

the riparian sites with near pristine housing more species

for B. stellatifolium and heavily invaded sites for A.

mearnsii, than restored sites (Table 2).

Generalised linear models indicated that for all arthro-

pods from two native tree taxa combined, near pristine sites

Table 2 Number of collected arthropod species (Sobs) and individ-

uals as well as the estimated number of species (Chao2 = second

order Chao estimator) from three tree species at sites that differ in

invasive status (near pristine, heavily invaded and restored)

Site Sobs Individuals Chao2 (±SD)

Overall 967 29,811 1215 (34.1)

Near pristine 868 9798 995.3 (44.6)

B. stellatifolium 479 5234 679.5 (32.3)

M. angustifolia 340 2373 609.3 (59.3)

A. mearnsii 346 2191 488.9 (31.8)

Heavily invaded 550 7667 666.9 (30.8)


M. angustifolia 295 2012 416.5 (30.3)

A. mearnsii 257 1349 510.8 (65.1)

Restored 615 12,346 857.9 (45.8)


M. angustifolia 297 2020 435.3 (33.0)

A. mearnsii 338 2350 486.9 (32.4)


123

had significantly higher alpha-diversity than heavily inva-

ded sites with intermediate alpha-diversity at restored sites

(F[2,6] = 72.9, p\ 0.001; Table 3). This was also true for

all host trees separately (F[2,42] = 72.02, p\ 0.001 for B.

stellatifolium; F[2,42] = 77.1, p\ 0.001 for M. angustifo-

lia; F[2,42] = 182.8, p\ 0.001 for A. mearnsii; Table 3).

Species richness (for all native tree species combined) was

highest at near pristine sites for most arthropod orders

(excluding the Ants and Lepidoptera), followed by restored

sites, with heavily invaded sites usually containing the least

number of species (Table 3). However, for most orders the

differences in alpha-diversity among degrees of plant

invasions for individual tree species were not significant.

The ants were more species rich at restored sites, but only

significantly so for all native tree taxa combined

(F[2,6] = 1.96, p = 0.05) and for B. stellatifolium

(F[2,42] = 2.37, p = 0.05; Table 3). Araneae alpha-diver-

sity was significantly lower at heavily invaded sites for

those associated with B. stellatifolium (F[2,6] = 3.19,

p\ 0.01; Table 3).

Arthropod species turnover among sites (b1)

When combining all arthropods collected on native hosts,

near pristine and restored sites which were statistically

similar had significantly higher b1-diversity (species turn-

over among sites) than heavily invaded assemblages

(F[2,6] = 8.91, p = 0.004; Table 4). However, the influ-

ence of plant invasions of riparian zones on b1-diversity

was non-significant for most arthropod taxa separately

except Coleoptera (F[2,6] = 9.31, p = 0.003; Table 4).

Coleopteran b1-diversity was significantly lower at

restored sites than at near pristine and heavily invaded

sites. b1-diversity for the different orders associated with

specific tree species varied little among sites with differing

plant invasions (Table 4).

Arthropod assemblage composition among sites

that differ in plant invasions (b2)

PERMANOVA analyses revealed that plant invasions of

riparian habitats significantly influenced arthropod assem-

blage composition when data from all native trees were

combined, with the exception of the Araneae, Coleoptera,

Diptera and Lepidoptera (Table 5; Fig. 2a). However,

nearly all pair-wise comparisons among sites for each

arthropod taxon separately (combined tree species data) did

not differ significantly except for the Hemiptera, Hyme-

noptera, Ants and Orthoptera (Table 5).

Riparian plant invasions significantly influenced overall

assemblage composition for B. stellatifolium and A.

mearnsii (Table 5). For B. stellatifolium, pair-wise com-

parisons indicated that the significant divergence between

communities at near pristine- and restored sites drove the

overall pattern (Table 5). For M. angustifolia, no signifi-

cant differences were detected in overall arthropod com-

munity assemblages but for a few taxa differences were

observed (Table 5). For A. mearnsii no differences were

found for pair-wise comparisons among the different plant

invasions, but overall plant invasions had a significant

influence on arthropod assemblages. Pair-wise comparisons

between arthropods from restored and heavily invaded

habitats never differed significantly, but comparisons

between near pristine and restored, and near pristine and

heavily invaded habitats did differ in a few cases (Table 5).

Comparisons of sites of the same plant invasion status for

Table 3 Summary results for

generalised linear models

(Poisson distribution and log-

link function) on species

richness data for the overall, and

eight most species-rich and

abundant taxonomic groups

Dependent variable Overall Tree species

B. stellatifolium M. angustifolia A. mearnsii

Species richness

Overall NP[R[HI NP[R = HI NP = R C HI NP = R C HI

Araneae NP[R[HI NP[R[HI NP = R = HI NP = R = HI

Coleoptera NP[R = HI NP[R = HI NP = R = HI NP = R = HI

Diptera NP[R[HI NP = HI C R NP = HI = R NP = HI = R

Hemiptera NP[R = HI NP[HI = R NP = R = HI NP = R = HI

Hymenopteraa NP[R = HI NP[R = HI NP = R = HI R = NP = HI

Lepidoptera NP = HI = R NP = R = HI HI = R = NP R = NP = HI

Ants R[NP = HI R[NP = HI R = NP = HI R = NP = HI

Orthoptera NP = R = HI NP = R = HI NP = R = HI NP = R = HI

Sites are ordered with those with the highest means on the left and the lowest on the right

NP near pristine, HI heavily invaded, R restored riparian habitat typesa All members of Hymenoptera except the Ants

= Signifies no significant differences,[signifies that habitats to the left are significantly more species-rich;

C signifies that the first habitat is significantly more species-rich than the last habitat


123

individual host tree species also indicated that plant inva-

sions of riparian habitats significantly influenced arthropod

assemblage composition, except for a few taxonomic

groups (Table 6).

When considering all arthropods from all two native

host tree species combined, the PCO plot showed sites

grouped strongly according to plant invasion (Fig. 2a).

This was also evident when considering arthropods col-

lected from the two native tree taxa respectively, but less so

when considering the arthropod communities associated

with A. mearnsii (Fig. 2d). Near pristine sampling sites for

B. stellatifolium were more closely grouped than heavily

invaded and restored sampling units that were more inter-

mixed (Fig. 2b). For M. angustifolia, heavily invaded units

separated out with near pristine and restored sites inter-

mixed (Fig. 2c). When considering collection of sites,

samples from specific sites tended to group together for all

three host trees (Fig. 2a).

Considering all arthropods collected for native hosts,

near pristine sites had proportionately higher numbers of

unique species 295 (27.8 %), higher than either the restored

214 (20.1 %) or heavily invaded 178 (16.7 %) riparian

habitats (Table 7). This was true for all separate tree spe-

cies. Araneae had proportionately higher numbers of unique

species 50 (4.7 %) in restored sites, higher than either the

near pristine 37 (3.5 %) or heavily invaded 28 (2.6 %)

riparian habitats (Table 7) for all separate tree species.

Discussion

Many studies have investigated the effects of IAPs on

species richness of arthropods. Although some report no

effect on certain arthropod taxa (e.g. Robertson et al.

2011), the vast majority indicate that IAPs have a negative

effect on arthropod taxa (e.g. Samways and Moore 1991;

Bultman and Dewitt 2008; Samways et al. 2011; Roets and

Pryke 2013). The variously invaded riparian habitats

compared in this study were found to differ in alpha

diversity of arthropods. Near pristine sites had higher

species richness than restored sites, with heavily invaded

sites housing fewest species for various arthropod taxa

except Ants. After removing IAPs these riparian habitats

can be recolonised by arthropods, with alpha diversity

returning to near pristine levels. Low arthropod species

richness in heavily invaded sites was expected given sim-

ilar results from other studies that have investigated the

impacts of IAPs on arthropod populations and communities

across a wide variety of habitats; both within South Africa

(Samways and Moore 1991; Samways et al. 2011; Roets

and Pryke 2013) and elsewhere (Toft et al. 2001; Bultman

and Dewitt 2008). However, unlike these studies ours

focused on arthropods associated with particular trees

rather than arthropods associated with the entire ecosystem.

Loss in some arthropod species in invaded sites may

therefore be independent of changes in plant diversity,

vegetation structure and microclimatic conditions (see Litt

et al. 2014). These changes would be worth exploring in

future studies.

No significant change in species richness of Ants was

detected between near pristine and heavily invaded sites.

Similarly, French and Major (2001) found no significant

differences in the species richness of Ants between areas of

South African Fynbos invaded by Acacia saligna and

native sites. In contrast to invaded sites, restored sites

supported significantly higher species richness of Ants.

This suggests that restored sites appear to be benefiting

ants, although the mechanisms behind this pattern are

Table 4 Results of tests for b1-

diversity for host trees using the

Jaccard resemblance measure

for each tree species using

different taxonomic groups

Plant invasions Tree species

B. stellatifolium M. angustifolia A. mearnsii

All NP = R C HI HI = R = NP NP = R = HI NP = R = HI

Araneae HI = R = NP HI = R = NP R = HI = NP R = NP = HI

Coleoptera NP = HI[R HI = NP = R NP = HI = R NP = R = HI

Diptera HI = R = NP HI = R = NP HI = R = NP HI = NP = R

Hemiptera R = HI = NP HI = R = NP R = NP = HI HI = R = NP

Hymenopteraa HI = R = NP HI = R = NP R = NP = HI NP = R = HI

Lepidoptera NP = R = HI HI = NP = R NP = R = HI HI = R = NP

Ants R = NP = R R = NP = HI R = NP = HI HI = NP = R

Orthoptera NP = R = HI NP = R = HI NP = HI = R R = HI = NP

Sites are ordered with those with the highest means on the left and the lowest on the right

NP near pristine, HI heavily invaded, R restored riparian habitat typesa All members of Hymenoptera except the Ants

= Signifies no significant differences,[signifies that habitats to the left are significantly more species-rich/

abundant; C signifies that the first habitat is significantly more species-rich/abundant than the last habitat


123

unclear. The reduced richness of Araneae in heavily

invaded sites could imply reduced predation pressure on

folivorous insects (members of Hemiptera and Coleoptera)

(Simao et al. 2010), eventually exacerbating folivore

damage to native plant species (Halaj and Wise 2001). This

decline in Araneae richness in heavily invaded sites

Table 5 Arthropod assemblage

beta-diversity (b2) from

PERMANOVA to determine

similarity in the composition of

arthropod assemblages among

riparian habitats that differ in

plant invasions for three tree

species and for the eight most

species-rich and abundant

taxonomic groups

Plant invasions NP versus HI NP versus R HI versus R

All 1.35* 1.33 1.11 1.04

Araneae 1.11 1.21 1.02 0.92

Coleoptera 1.2 1.23 1.31 0.71

Diptera 1.01 1.12 1.17 0.68

Hemiptera 2.25** 1.48* 1.46* 1.56

Hymenopteraa 2.02** 1.51* 1.46* 1.31

Lepidoptera 0.74 0.84 0.84 0.89

Ants 2.09* 1.80* 1.48* 0.95

Orthoptera 2.89** 1.94* 1.64* 1.47*

B. stellatifolium

Overall 1.69** 1.29 1.4* 1.17

Araneae 1.18 1.15 1.09 1.01

Coleoptera 2.44** 1.53 1.79* 1.33

Diptera 0.93 1.06 1.01 0.81

Hemiptera 1.76** 1.37 1.29 1.31

Hymenopteraa 1.80** 1.36 1.36 1.31

Lepidoptera 0.94 1.04 0.89 0.95

Ants 1.27 1.65 0.63 1.08

Orthoptera 1.74 1.64 1.16 1.12

M. angustifolia

Overall 1.41 1.17 0.99 1.39

Araneae 1.42* 1.27 1.21 1.09

Coleoptera 1.77** 1.44* 1.35* 1.21

Diptera 1.33* 1.27 1.31 0.85

Hemiptera 1.72** 1.24 1.43* 1.26

Hymenopteraa 1.65 1.44 1.22 1.18

Lepidoptera 1.16 1.02 1.05 1.16

Ants 1.91 1.69 1.35 1.08

Orthoptera 2.20* 1.52 1.92 0.82

A. mearnsii

Overall 1.57* 1.37 1.25 1.13

Araneae 1.18 1.05 1.10 1.09

Coleoptera 1.44* 1.23 1.24 1.12

Diptera 1.36 1.14 1.34 1.01

Hemiptera 1.36* 1.23 1.11 1.15

Hymenopteraa 1.58* 1.35 1.27 1.14

Lepidoptera 1.00 0.96 1.28 0.75

Ants 1.40 1.08 1.35 1.13

Orthoptera 1.87* 1.62 1.54 0.93

Figures represent F-(second column) and t-values (column 3–5), df = 8, number of permutations for each

analysis = 9999

NP near pristine, HI heavily invaded, R restored riparian habitats

* P\ 0.05; ** P\ 0.01a All members of Hymenoptera with the exception of Ants


123

suggests that it would be beneficial to quantify damage

levels to plants across all plant invasions to explore the

possible consequences of altered Araneae richness.

In contrast to Araneae species richness, Lepidoptera and

Orthoptera, richness were not affected by plant invasions.

This suggests that current management practices in riparian

zones in South Africa are not having a major impact on

their species richness and that these orders may be less

important as indicator groups when assessing IAP status.

These findings are similar to Harris et al.’s (2004) argu-

ment that invasive plants do not necessarily have to impact

biodiversity negatively. In their study, Ulex europaeus (an

exotic invasive shrub in New Zealand) supported more

insect species of some taxonomic groups than did native

Kanuka trees (Kunzea ericoides).

Considering all arthropods together, restored and near

pristine sites had much more homogenous arthropod

communities as compared to heavily invaded sites, based

Fig. 2 Principal Coordinate Ordination (PCO) plots of arthropod

assemblages from near pristine (circle), heavily invaded (star), and

restored (triangle) riparian habitats for a all arthropods from all host

trees combined, b arthropods collected from, B. stellatifolium,

c arthropods collected from M. angustifolia and d arthropods

collected from A. mearnsii. The ellipses represent sampling units

which were 25 % similar. UD Upper Dwars, DK Du toits Kloof, BK

Bains Kloof, MD Mid Dwars, LW Lower Wit, MW Mid Wit, LD

Lower Dwars, DT Du Toit, and UM Upper Molenaars collection sites


123

on PERMDISP results. This suggests that after restoration

of a riparian ecosystem, a site is usually recolonised by a

community consisting of similar, abundant arthropod taxa.

It is possible that, given enough time, rarer arthropod taxa

would also recolonise the restored habitats and ultimately

increase variability between these areas. Possible reasons

for significantly higher b1-diversity for arthropods in near

pristine as compared to heavily invaded sites are numerous,

but may include: (1) higher heterogeneity in both plant

species composition and structure (Walz 2011); there is

current evidence that diverse habitats support higher bio-

logical diversity than monotypic ones, thus allowing more

Table 6 Main test of arthropod assemblage beta-diversity (b2) from

PERMANOVA to determine similarity in the composition of

arthropod assemblages among riparian sites that are similar in plant

invasions for three tree species and for the eight most species-rich and

abundant taxonomic groups

Near pristine Heavily invaded Restored

B. stellatifolium

Overall 3.64*** 3.39*** 3.22***

Araneae 2.87*** 1.58* 2.83***

Coleoptera 4.22*** 3.19*** 4.16***

Diptera 3.83*** 2.24*** 2.89***

Hemiptera 2.61*** 2.43*** 1.51*

Hymenopteraa 3.45*** 3.40*** 3.22***

Lepidoptera 1.74* 1.35 1.66*

Ants 1.98* 1.81* 1.81*

Orthoptera 1.75* 1.49 0.87

M. angustifolia

Overall 3.24*** 2.91*** 4.07***

Araneae 2.12** 1.98*** 3.59***

Coleoptera 2.97*** 3.48*** 3.03***

Diptera 2.72*** 2.28*** 3.42***

Hemiptera 3.68*** 2.17*** 4.02***

Hymenopteraa 2.45** 4.21*** 3.13***

Lepidoptera 1.74* 1.57 2.21**

Ants 1.21 1.07 5.74***

Orthoptera 1.03 1.26 1.72

A. mearnsii

Overall 3.63*** 3.89*** 3.39***

Araneae 1.60** 4.85*** 1.68***

Coleoptera 6.45*** 3.00*** 3.44***

Diptera 4.73*** 4.02*** 4.46***

Hemiptera 2.48*** 3.14*** 3.31***

Hymenopteraa 4.52*** 3.02*** 2.81***

Lepidoptera 1.83* 1.88** 1.39

Ants 0.95 2.15** 1.38

Orthoptera 1.10 1.61 1.07

Figures represent F-values, df = 44, number of permutations for each

analysis = 9999

* P\ 0.05; ** P\ 0.01; *** P B 0.001a All members of Hymenoptera with the exception of Ants

Table 7 Number of common arthropod species (considering only

those with more than four individuals collected throughout the study

period) that were unique to a specific habitat or tree species, for

various assemblages collected from CFR riparian habitats (percentage

of total in parenthesis)

Near pristine Heavily invaded Restored

All 295 (27.8) 178 (16.7) 214 (20.1)

Araneae 37 (3.5) 28 (2.6) 50 (4.7)

Coleoptera 59 (6.5) 31 (2.9) 41 (3.9)

Diptera 42 (3.9) 19 (1.8) 23 (2.2)

Hemiptera 69 (6.5) 29 (2.7) 33 (3.1)

Hymenopteraa 28 (2.6) 17 (1.6) 19 (1.8)

Lepidoptera 12 (1.1) 8 (0.8) 11 (1.0)

Ants 4 (0.4) 2 (0.2) 5 (0.5)

Orthoptera 8 (0.8) 3 (0.3) 6 (0.6)

B. stellatifolium

Overall 270 (39.4) 188 (27.4) 219 (31.9)

Araneae 60 (8.8) 27 (3.9) 46 (6.7)

Coleoptera 58 (8.5) 50 (7.3) 45 (6.6)

Diptera 42 (6.1) 31 (4.5) 29 (4.2)

Hemiptera 57 (8.3) 32 (4.6) 40 (5.8)

Hymenopteraa 22 (3.2) 16 (2.3) 20 (2.9)

Lepidoptera 7 (1.0) 8 (1.2) 10 (1.5)

Ants 4 (0.6) 3 (0.4) 7 (1.0)

Orthoptera 3 (0.4) 2 (0.3) 1 (0.1)

M. angustifolia

Overall 241 (38.9) 199 (32.1) 220 (35.5)

Araneae 49 (7.9) 32 (5.2) 54 (8.7)

Coleoptera 47 (7.6) 49 (7.9) 52 (8.4)

Diptera 34 (5.5) 39 (6.3) 29 (4.7)

Hemiptera 56 (9.0) 36 (5.8) 35 (5.7)

Hymenopteraa 32 (5.2) 22 (3.6) 23 (3.7)

Lepidoptera 4 (0.6) 6 (1.0) 3 (0.5)

Ants 3 (0.5) 1 (0.2) 5 (0.8)

Orthoptera 0 (0.0) 2 (0.3) 4 (0.6)

A. mearnsii

Overall 256 (38.7) 187 (28.3) 262 (39.6)

Araneae 49 (7.4) 31 (4.7) 54 (8.2)

Coleoptera 53 (8.0) 35 (5.3) 59 (8.9)

Diptera 35 (5.3) 31 (4.7) 28 (4.2)

Hemiptera 49 (7.4) 30 (4.5) 40 (6.1)

Hymenopteraa 29 (4.4) 25 (3.8) 29 (4.4)

Lepidoptera 5 (0.8) 11 (1.7) 13 (1.9)

Ants 7 (1.1) 6 (0.9) 3 (0.5)

Orthoptera 2 (0.3) 4 (0.6) 5 (0.8)

a All members of Hymenoptera with the exception of Ants


123

species to coexist (Mlambo et al. 2011), (2) spatial auto-

correlation i.e. sites that are further apart have a tendency

to differ drastically in arthropod species composition

(Horak 2013).

Not all arthropods associated with native host taxa

reacted similarly to plant invasions and restoration. For

example, b2-diversity of arthropod communities on B.

stellatifolium were fairly similar between restored and

invaded sites, while on M. angustifolia the arthropod

communities from restored sites were more similar to near

pristine sites. Restoration success therefore varies consid-

erably when considering the trends associated with indi-

vidual plant taxa and their respective arthropod

communities and they need different lengths of time to

regenerate. The reason for this is unclear, but may be due to

changes in plant characteristics (e.g. physical structure, leaf

chemistry, and host abundance) associated with plant

invasion (for example see Lathrop et al. 2003). It is pos-

sible that M. angustifolia characteristics that may alter

quality of habitats for arthropods did not change in the

presence of IAPs, hence arthropods were quick to recolo-

nise M. angustifolia individuals after removal of IAPs.

Conversely, IAPs appear to have heavily influenced the

characteristics of B. stellatifolium thereby delaying the

return of arthropod communities to their original state.

When considering A. mearnsii, arthropod communities,

b2-diversity varied substantially among different collection

sites and among plant invasions. However, plant invasion

status of sampling sites had a lesser effect on grouping of

communities than it did on collection sites. This can be

explained if one considers that most arthropods associated

with A. mearnsii are actually associated with the sur-

rounding vegetation rather than A. mearnsii itself (as can

be expected from a non-native plant in accordance with the

enemy release hypothesis (Wolfe 2002; Siemann and

Rogers 2003; van der Colff et al. 2015). Although limited

information exists on the arthropod communities of A.

mearnsii in its invaded range, it is colonised almost

exclusively by native arthropods within forestry plantations

(Govender 2007; DEA 2009). The arthropod communities

associated with A. mearnsii are therefore expected to reflect

the general communities associated with the specific sites

where it is found.

The results of this study are largely in accordance with

other studies (e.g. Wishart et al. 2002; Samways et al.

2011) that found that individual rivers of Fynbos biore-

gions of the Western Cape have specific arthropod com-

munities (i.e. catchment signatures). This is not surprising,

given the high spatial variability in Mediterranean-type

ecosystems (Caterino 2007). Interestingly, the three Fynbos

studies mentioned were limited to aquatic invertebrates

while this study focused on terrestrial invertebrates. Thus,

the phenomenon of specific river catchment arthropod

communities prevails even when the organisms in these

systems are not directly dependant on the water itself.

Results further highlight the importance of conserving

and maintaining near pristine sites for sustaining overall

diversity in riparian habitats as these contain numerous

unique species (particularly Hemiptera and Coleoptera).

Unique species are perceived as important in ecological

systems and their preservation is often the ultimate aim of

biological monitoring (Lenat and Resh 2001). The

recolonisation of restored habitats by particularly rare

arthropods will also depend greatly on the availability of

nearby suitable habitat. It has previously been demon-

strated that fragments of natural habitat in CFR are

important for the conservation of many endemic species

(Kemper et al. 1999).

To conclude, the above results clearly underscore that

alpha and beta-diversity of arthropods are greatly impacted

by different plant invasions of riparian habitats. Removal of

IAPs appears to benefit species richness of the majority of

taxonomic groups. Arthropod beta diversity demonstrated

that a change in species composition may be a better

measure than alpha diversity to detect shifts in arthropod

communities induced by different plant invasion levels of

riparian habitats than species richness alone (e.g. Pryke

et al. 2013). These changes in community composition may

have profound influences on the normal functioning of

riparian ecosystems. Restoration success should also be

evaluated on a per species basis when considering arthro-

pods associated with foliage as recovery of arthropods on

different hosts appears to vary between host species.

Acknowledgments Thanks to Marno Fourie and Oliver Kambaj for

field assistance and reviewers for helpful suggestions to improve this

manuscript. We thank private landowners and CapeNature for access

to field sites (Permit Number: AAA008-00027-0028). This study was

funded by the DST/NRF Centre of Excellence in Tree Health

Biotechnology.

Compliance with ethical standards

Conflict of interest The authors declare that they have no conflict

of interest.

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Restoration of invaded Cape Floristic Region riparian